5. POTENTIAL FOR HUMAN EXPOSURE · 5. POTENTIAL FOR HUMAN EXPOSURE 5.1 OVERVIEW Chlorinated dioxins (CDDs) are a family of compounds that includes some extremely toxic and potent

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5. POTENTIAL FOR HUMAN EXPOSURE

5.1 OVERVIEW

Chlorinated dioxins (CDDs) are a family of compounds that includes some extremely toxic and potent

congeners. The two most toxic of the CDDs in mammals are 2,3,7,8-TCDD and 1,2,3,7,8-PeCDD (Buser

1987; Poland and Knutson 1982; Safe 1986; WHO 1997). In general, the more toxic congeners to mammals

appear to be the 2,3,7,8-substituted tetra-, penta-, and hexachloro- compounds, (e.g., 2,3,7,8-TCDD,

1,2,3,7,8-PeCDD, 1,2,3,4,7,8-HxCDD, 1,2,3,6,7,8-HxCDD, and 1,2,3,7,8,9-HxCDD) (Poland and Knutson

1982; Safe 1986; WHO 1997). A more detailed discussion of the relative toxicities of the different CDD

congeners is given in Section 2.5, Relevance to Public Health.

CDDs usually occur in the environment concurrently with other chemicals such as chlorinated dibenzo

furans (CDFs). CDDs and CDFs are highly persistent compounds and have been detected in air, water,

soil, sediments, animals, and foods. CDFs include 135 congeners, which are structurally similar to CDDs

and which elicit a number of similar toxicological and biochemical responses in animals (for more

information on CDFs see ATSDR 1994). CDDs and CDFs are released to the environment during

combustion processes (e.g., municipal solid waste, medical waste, and industrial hazardous waste

incineration, and fossil fuel and wood combustion); during the production, use, and disposal of certain

chemicals (e.g., PCBs, chlorinated benzenes, chlorinated pesticides); during the production of bleached

pulp by pulp and paper mills; and during the production and recycling of several metals (Buser et al.

1985; Czuczwa and Hites 1986a, 1986b; Oehme et al. 1987, 1989; Zook and Rappe 1994). The EPA

has developed procedures for estimating risks associated with exposures to mixtures of CDDs and CDFs

in environmental matrices (EPA 1989e). This approach is based on the assignment of 2,3,7,8-TCDD

toxic equivalence factors (TEFs) to CDD/CDF congeners or homologues in complex mixtures. The

rationale behind the use of TEFs is explained in Section 2.5, Relevance to Public Health. Although the

focus of this profile is CDDs, it should be recognized that most exposure scenarios involve exposure

to CDDs, CDFs, and the non-ortho polychlorinated biphenyls (PCBs) that have CDD-like toxicity;

many of these exposure scenarios are discussed in this chapter. These exposures are usually reported

in TEQs (for more information see Section 2.5, Relevance to Public Health, Toxic Equivalency Factors

[TEFs] and Toxic Equivalents [TEQs]). Over the past several years sets of TEFs have been

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developed, varying slightly from one to another. The reader is encouraged to consult the original literature

for specific details on TEQs computation.

CDDs (TCDD, PeCDD, HxCDD, HpCDD, OCDD) are ubiquitous in the environment (Podoll et al. 1986).

Although all of the sources or processes that contribute to CDDs in the environment have not been

identified, CDDs are known to be formed in the manufacture of chlorinated intermediates and pesticides,

during smelting of metals (EPA 1998j), in the incineration of municipal, medical, and industrial wastes

(Podoll et al. 1986), and from the production of bleached wood pulp and paper (Fletcher and McKay 1993).

CDDs are also found in emissions from the combustion of various other sources, including coal-fired or oil-

fired power plants, wood burning, and home heating systems (Chiu et al. 1983; Czuczwa and Hites 1984;

EPA 1998j; Gizzi et al. 1982; Thoma 1988). Generally, the more highly chlorinated CDDs are the most

abundant congeners present in the emissions from these combustion sources. CDDs also occur in other

combustion products (e.g., cigarette smoke) (Bumb et al. 1980; Lofroth and Zebuhr 1992; Muto and

Takizawa 1989), automobile exhaust from cars running on leaded gasoline with chlorine scavengers and to

a lesser extent from cars running on unleaded gasoline (Bingham et al. 1989; Marklund et al. 1987, 1990),

and diesel exhaust (Jones 1995; Cirnies-Ross et al. 1996). CDDs/CDFs can form during the synthesis and

combustion of chlorine-containing materials, such as polyvinylchloride (PVC), in the presence of naturally

occurring phenols, vegetation treated with phenoxy acetic acid herbicides, paper and wood treated with

chlorophenols, and pesticide-treated wastes (Arthur and Frea 1989).

CDDs occur as contaminants in the manufacture of various pesticides and, as a result, have been released

to the environment during use of these pesticides. 2,3,7,8-TCDD is a by-product formed in the manu

facture of 2,4,5-trichlorophenol (2,4,5-TCP) (Arthur and Frea 1989). 2,4,5-TCP was used to produce the

bactericide, hexachlorophene, and the chlorophenoxyherbicide, 2,4,5-trichlorophenoxy acid (2,4,5-T).

Trichlorophenol-based herbicides have been used extensively for weed control on crops, rangelands,

roadways, right-of-ways, etc. Various formulations of 2,4-dichlorophenoxy acetic acid (2,4-D)

contaminated mainly with higher chlorinated CDDs/CDFs and 2,4,5-T contaminated mainly with

2,3,7,8-TCDD were used extensively for defoliation and crop destruction by the American military during

the Vietnam War. Although six herbicides were used (Orange, Purple, Pink, Green, White, and Blue),

herbicide Orange (Agent Orange) was the primary defoliant (Wolf et al. 1985). Hexachlorophene use has

been restricted by the FDA and its disposal is regulated by EPA under the Resource Conservation and

Recovery Act (RCRA). In 1983, EPA canceled registration for all chlorophenoxy herbicides used on

foods, rice paddies, pastures, and rangelands (IARC 1986b). 2,4,5-T can no longer be used legally in the

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United States for any purpose (IARC 1986b). Other countries, including Canada, Sweden, the Netherlands,

Australia, Italy, and the Federal Republic of Germany, have also canceled registrations for 2,4,5-T (IARC

1986b), but many other countries have not. Currently, 2,4,5-T can be produced with lower 2,3,7,8-TCDD

concentrations than were previously possible. 2,4,5-TCP production has been discontinued in many

countries, including the United States, Canada, the United Kingdom, the Federal Republic of Germany,

and Austria (IARC 1986a). HxCDD, HpCDD, and OCDD are known contaminants of pentachlorophenol

(PCP), primarily a wood preservative and pesticide, which was used extensively in the 1970s and is still

used today (to a lesser extent) in the lumber industry. PCP is currently registered as a restricted-use

pesticide in the United States (Sine 1990).

Although little definitive data exist to prove or disprove that CDDs form during natural processes, results

from dated sediment cores have shown that there were significant increases in CDDs and CDFs after about

1940 (Czuczwa and Hites 1984, 1986b, 1986b) and lower levels of CDDs are currently found in persons

from less industrialized countries (Schecter et al.1991a). The congener/homologue profile of the

sediments was similar to that of atmospheric samples, strongly suggesting that combustion processes were

the source of CDDs in the sediments. The historical increase in CDDs/CDFs also was similar to the trends

for the production, use, and disposal of chlorinated organics, suggesting that accumulation of these

compounds in the environment is a recent phenomenon related to the production, use, and subsequent

incineration of chlorinated organic chemicals (Schecter et al. 1988).

CDDs are ubiquitous in the environment and are found at low background levels (parts per trillion [ppt] or

parts per quadrillion [ppq]) in the air, water, and soil. Lower levels are found in biological and environ

mental samples from less industrialized rural regions than in those from more industrialized urban regions

(Czuczwa and Hites 1986a; Des Rosiers 1987; Edgerton et al. 1989; Schecter et al. 1989e, 1989g, 1991a,

1994d; Tiernan et al. 1989b). HpCDD and OCDD are the most common CDDs found in environmental

samples (Christmann et al. 1989b; Clement et al. 1985, 1989; Pereira et al. 1985; Reed et al. 1990; Tashiro

et al. 1989a; Tiernan et al. 1989b).

The environmental fate and transport of CDDs involve volatilization, long-range transport, wet and dry

deposition, photolysis, bioaccumulation, and biodegradation (Kieatiwong et al. 1990). CDDs strongly

partition to soils and sediments. Due to their low vapor pressure and low aqueous solubility and their

strong sorption to particulates, CDDs are generally immobile in soils and sediments. Although most

biological and nonbiological transformation processes are slow, photolysis has been shown to be relatively

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rapid. Photolysis is probably the most important transformation process in environmental systems into

which sunlight can penetrate (Kieatiwong et al. 1990). Estimates of the half-life of 2,3,7,8-TCDD on the

soil surface range from 9 to 15 years, whereas the half-life in subsurface soil may range from 25 to

100 years (Paustenbach et al. 1992). CDDs have been shown to bioaccumulate in both aquatic and

terrestrial biota. CDDs have a high affinity for lipids and, thus, will bioaccumulate to a greater extent in

organisms with a high fat content.

Over the past decade, typical concentrations of CDDs in urban air in the United States have averaged

2.3 pg/m3, with OCDD and HpCDD homologues predominating and 2,3,7,8-TCDD being the least

common congener (Smith et al. 1992). CDD concentrations range as follows: OCDD, 0.44–3.16 pg/m3;

HpCDD, 0.21–4.4 pg/m3; HxCDD, 0.6–0.63 pg/m3; PeCDD, not detected to 0.1 pg/m3; and 2,3,7,8-TCDD,

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and 1,2,3,7,8-PeCDD, the CDDs currently believed to be most toxic to vertebrates (WHO 1997), were

found in fish tissue at 70% and 54% of the sites, respectively. 2,3,7,8-TCDD was found at a mean

concentration of 6.9 ppt and a maximum concentration of 204 ppt, and 1,2,3,7,8-PeCDD was found at a

mean concentration of 2.38 ppt and a maximum concentration of 54 ppt. With respect to source

categories, fish collected near pulp and paper mills using chlorine had the highest median 2,3,7,8-TCDD

concentration (5.66 ppt), compared to the second highest median 2,3,7,8-TCDD concentrations of 1.82 ppt

at refinery/other industrial sites, and the third highest median 2,3,7,8- TCDD concentration of 1.27 ppt at

Superfund sites. Similarly, with respect to source categories, fish collected near pulp and paper mills using

chlorine had the highest median 1,2,3,7,8-PeCDD concentration (1.52 ppt), compared to the second

highest median concentrations of 1.35 ppt at refinery/other industrial sites, and the third highest median

concentration of 1.09 ppt at industrial/urban sites.

The detection of CDDs in blood, adipose tissue, breast milk, and other tissue samples from the general

population indicates universal exposure to CDDs from environmental sources (Fürst et al. 1994; Orban et

al. 1994; Patterson et al. 1986a; Ryan et al. 1986, 1993a; Schecter and Gasiewicz 1987a, 1987b; Schecter

et al. 1986b, 1989e; Stanley 1986; Stanley et al. 1986). The general population is exposed to CDDs

released from industrial and municipal incineration processes; exhausts from automobiles using leaded

gasoline; cigarette smoke; and foods, including human milk (Pohl and Hibbs 1996; Schecter et al. 1994e).

The major source (>90%) of exposure for the general population, however, is primarily associated with

meat, dairy products, and fish (Beck et al. 1989a; Schaum et al. 1994; Schecter et al. 1994d, 1994e,

1996a). CDDs are transferred through the placenta to the fetus, by breast milk to infants and young

children, and by lifelong dietary ingestion. Workers involved with incineration operations or those who

have been or may be involved in the production, use, or disposal of trichlorophenol, phenoxyherbicides,

hexachlorophene, pentachlorophenol and other compounds that contain impurities of CDDs are at a greater

risk from exposure to CDDs and TEQs (Päpke et al. 1992; Schecter and Ryan 1988; Schecter et al. 1991).

Individuals in the general population who may be exposed to potentially higher levels of CDDs include

recreational and subsistence fishers (including many native Americans) and their families living in CDD-

contaminated areas who consume large quantities of fish from contaminated waters (CRITFC 1994; Ebert

et al. 1996), subsistence hunters such as the Inuit of Alaska who consume large quantities of wild game

(particularly marine mammals) (Dewailly et al. 1993; Hebert et al. 1996; Norstrom et al. 1990),

subsistence farmers and their families living in areas contaminated with CDDs who consume their own

farm-raised beef and dairy products (EPA 1996b; McLachlan et al. 1994), individuals who live in the

vicinity of an industrial or municipal incinerator, or individuals who live in the vicinity of the

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126 hazardous waste sites where CDDs (and more especially where 2,3,7,8-substituted CDDs) have been

detected (Gough 1991; Liem et al. 1991; Pohl et al. 1995; Riss et al. 1990; Wuthe et al. 1993).

2,3,7,8-TCDD has been identified in at least 91 of 1,467 current or former EPA National Priorities List

(NPL) hazardous waste sites (HazDat 1998). However, the number of sites evaluated for 2,3,7,8,-TCDD is

not known. The frequency of these sites within the United States can be seen in Figure 5-1. Of these sites,

90 are located in the United States and 1 is located in the Commonwealth of Puerto Rico (not shown).

Total CDDs (including TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD) have been identified in 126,

105, 34, 43, 49, and 53 sites, respectively, of the 1,467 hazardous waste sites on the NPL. The frequency

of these sites within the United States for total CDDS, TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD,

respectively, can be seen in Figures 5-2 through 5-7. Of the 126 sites with total CDD detections, 125 are

located in the United States and 1 site is located in the Commonwealth of Puerto Rico (not shown). Of the

105 sites with total TCDD detections, 104 are located in the United States and 1 site is located in the

Commonwealth of Puerto Rico (not shown). Of the sites with PeCDD, HxCDD, HpCDD, and OCDD

detections, all 34, 43, 49, and 53 sites, respectively, are located in the United States.

5.2 RELEASES TO THE ENVIRONMENT

CDDs have been measured in all environmental media including ambient air, surface water, groundwater,

soil, and sediment. While the manufacture and use of chlorinated compounds, such as chlorophenols and

chlorinated phenoxy herbicides, were important sources of CDDs to the environment in the past, the

restricted manufacture of many of these compounds has substantially reduced their current contribution to

environmental releases. It is now believed that incineration/combustion processes are the most important

sources of CDDs to the environment (Zook and Rappe 1994). Important incineration/combustion sources

include: medical waste, municipal solid waste, hazardous waste, and sewage sludge incineration; industrial

coal, oil, and wood burning; secondary metal smelting, cement kilns, diesel fuel combustion, and

residential oil and wood burning (Clement et al. 1985; Thoma 1988; Zook and Rappe 1994).

Emissions to the atmosphere from incineration and combustion sources result in the wide-spread

distribution of CDDs. Consequently, CDDs are found at low levels in rural soils as well as in sediments of

otherwise pristine waterbodies. Much of the CDD deposits from wet and dry deposition ultimately

become components of urban runoff which enter rivers, streams, and estuaries directly or through

stormwater outfalls and combined sewer overflows (CSOs). In a recent study, Huntley et al. (1997) used

statistical

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pattern matching techniques (principal components analysis) to evaluate CDD congener patterns in

sediment samples collected adjacent to several CSOs. According to these authors, the presence of these

unique CDD/CDF congener patterns in sediment adjacent to CSOs suggested that these CSOs were a likely

source given the industrial, residential, and stormwater inputs to the combined sewer overflow system.

Such statistical techniques have been applied elsewhere to CDD congener pattern matching in an effort to

identify specific sources of CDDs. Wenning et al. (1993a, 1993b) also applied principal components

analysis to Newark Bay Estuary sediments and found that most of the congener fingerprint patterns were

related to combustion/incineration sources. More recently, Ehrlich et al. (1994) applied polytopic vector

analysis, a fingerprinting technique that “unmixes” the CDD/CDF patterns, and concluded that the primary

sources of CDD/CDFs in Newark Bay Estuary sediments were combustion/incineration, sewage-related

sources, and PCB-related sources. Statistical techniques that have proven useful for identifying sources of

CDDs have recently been reviewed (Wenning and Erickson 1994). Future efforts to reduce the release of

CDDs to the environment will require additional analysis of the distributional patterns of CDDs in

environmental media, which may also provide information on sources still to be identified.

5.2.1 Air

The key sources of CDD releases to air are from anthropogenic combustion processes and the production

and use of chemicals contaminated with CDDs. Some evidence suggests that natural combustion

processes (e.g., forest fires or volcanic activity) may also be sources of CDDs, but to a much smaller

extent. Toxics Release Inventory (TRI) data are not available for CDDs since CDD releases are not

required to be reported (EPA 1995g).

Combustion Processes. Combustion processes generate CDDs, CDFs, and other halogenated aromatic compounds (Czuczwa and Hites 1984, 1986a, 1986b). Most of the direct releases of CDDs and

CDFs from combustion processes are to the air (Czuczwa and Hites 1984, 1986a, 1986bc). CDDs and

CDFs may be found in particulates released from the combustion of most types of organic material and

limited evidence suggests that they may also result from trace chemical reactions in fire (Bumb et al. 1980;

Crummett 1982; Safe 1990). The processes involved in the formation of CDDs and CDFs consist of

numerous chemical reactions that occur during combustion of organic compounds in the presence of

chlorinated material. The EPA has recently identified stationary source categories that release

2,3,7,8-TCDD TEQ to the atmosphere (EPA 1998j). The percentage contribution of the five highest

source categories are: 68% from municipal waste incineration, 12.3% from medical waste incineration,

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8.9% from Portland cement manufacture hazardous waste kilns, 3.5% from secondary aluminum smelting,

and 3.0% from other biological incineration. These five source categories account for 95.9% of all

stationary emissions of 2,3,7,8-TCDD TEQ to the air.

The "Trace Chemistries of Fire Hypothesis" suggests that CDDs and CDFs can also form during a variety

of combustion processes including natural ones, such as forest fires and volcanic eruptions (Crummett

1982). However, there is very limited evidence suggesting that such natural processes could be minor

sources of these compounds in the environment. Only data from one study were found that directly

measured CDD/CDFs in actual emissions from forest fires. Tashiro et al. (1990) detected the concen

tration of total CDD/CDFs in air ranging from 15 to 400 pg/m3. The samples were collected from fixed

collectors 10 m above the ground and from aircraft flying through the smoke. Soil samples collected

before the burn detected 43 ppt of OCDD in 1 of 4 samples tested. After the burn, OCDD was detected in

3 of 4 soil samples at concentrations of 46, 100, and 270 ppt. Because the small sample size precluded

statistical analysis, no further conclusions were drawn by the authors. Thomas and Spiro (1995), however,

estimated that forest and agricultural burning accounted for the third largest emission of CDD/CDF in the

United States (30 kg/year), behind municipal waste incineration (200 kg/year) and hospital incinerators

(40 kg/year) although the inclusion of agricultural burning, which may include acreage treated with long-

lived organochlorine pesticides, may skew the values higher than would be expected from forest fires

alone. Failure to find CDDs in ancient mummies or ancient frozen Eskimo tissues is another indication

that the “Trace Chemistries of Fire Hypothesis” may have little bearing on human exposure (Ligon et al.

1989; Schecter et al. 1988; Tong et al. 1990). The EPA recently found elevated levels of 2,3,7,8-TCDD in

two chickens that were traced to clay (used as an anti-caking additive in soybean animal meal) derived

from clay deposits mined at the Kentucky-Tennesse Ball Clay Company in Crenshaw, Mississippi.

(Chemical Regulation Reporter 1997a, 1997b). However, no information on the origin of the 2,3,7,8

TCDD, either natural or anthropogenic, was presented.

The issue of natural sources of CDD/CDF is interesting, but historical deposition records strongly

implicate anthropogenic activity as the major source of CDD/CDFs (Thomas and Spiro 1996). These

authors further suggest that the historic record on CDD/CDF deposition provided by sediment cores

strongly implies that anthropogenic sources have been overwhelmingly dominant. Sediment cores from

Siskwit Lake on a remote island in northern Lake Superior, provide a historic record of atmospheric CDD

fluxes (Czuczwa and Hites 1986a). An 8-fold increase in the CDD/CDF deposition rate (from approx

imately 4–30 pg/cm2/year) occurred between 1940 and 1970, corresponding to a great expansion in the

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industrial use of chlorine (Thomas and Spiro 1996). The decrease in deposition rate of about 30% (from

30 to 24 pg/cm2 /year) from 1970 to the mid 1980s parallels decreased production and use of chlorophenols

(pesticide registrations for 2,4,5-T and Silvex were discontinued in 1983 and 1984, respectively) (IARC

1977; Sine 1990) and reductions in municipal incinerator emission resulting from improvements in design,

pollution controls, and operation of these facilities (Thomas and Spiro 1996). It is difficult to reconcile

these trends with predominantly natural sources, especially since the total area of U.S. forests consumed by

forest fires diminished by more than a factor of 4 between 1940 and 1970 through more effective fire

control (Thomas and Spiro 1996).

Although the production of CDDs during combustion processes are highlighted here, most samples from

combustion sources show a complex mixture of isomers and congeners of CDDs and CDFs which vary in

their relative concentrations (Kolenda et al. 1994; Nestrick and Lamparski 1983; Vikelsoe et al. 1994).

CDDs have been detected in emissions (flue gas and fly ash) from municipal, hazardous waste, and

industrial incinerators (Buser 1987; Oppelt 1991; Sedman and Esparza 1991; Schecter 1983). Combustion

of materials, such as vegetation treated with phenoxy acetic acid herbicides, paper and wood treated with

chlorophenols, pesticide-treated wastes, and polyvinylchloride (PVC) in the presence of naturally

occurring phenols, may lead to CDDs and CDD precursors (Arthur and Frea 1989). PVC is known to

yield a small amount of chlorobenzene upon pyrolysis, which in turn thermally decomposes to CDDs and

CDFs (Lustenhouwer et al. 1980). CDDs have also been detected in fly ash from an oil-fired power plant,

in city dust, in commercial sludge fertilizer, in particulate deposits in car and truck mufflers, in exhaust

from vehicles powered with leaded and unleaded gasoline and diesel fuel, in cigarette smoke, and in soot

from home fireplaces and from PCB and chlorinated benzene contaminated transformer fires (Bumb et al.

1980; Hutzinger et al. 1985; Lofroth and Zebuhr 1992; Marklund et al. 1987, 1990; Muto and Takizawa

1989; Schecter 1983; Thoma 1988). Dichloroethane, the chlorinated additive in leaded gasoline, is also a

source of CDDs (Marklund et al. 1987). The dichloroethane acts as a scavenger to prevent the deposition

of lead compounds in engines (Safe 1990). Although the data indicate that CDDs result from diverse

processes, the relative contributions of these sources and other unidentified sources to the presence of

CDDs in the atmosphere are not known.

A mixture of CDDs (TCDD, PeCDD, HxCDD, HpCDD, OCDD) has been found in emissions (both

particles and flue gases) from various combustion sources, including municipal incinerators, power plants,

wood burning, home heating systems, and petroleum refining (Chiu et al. 1983; Czuczwa and Hites 1984;

Gizzi et al. 1982; Nessel et al. 1991; Thoma 1988; Thompson et al. 1990). In individual samples of

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emissions from an urban incinerator, HxCDDs and OCDD were often the most abundant CDDs found,

although the homologue pattern can be quite variable (Gizzi et al. 1982). Emission of TCDD from

municipal waste combustion ranged from 0.018 ng/m3 to 62.5 ng/m3 depending on the type of combustion

facility (Roffman and Roffman 1991). A municipal solid waste incinerator sampled in 1988 contained an

average TCDD concentration of 0.0012 ng/m3, where OCDD was present at 1.2 ng/m3, and HxCDD was

present at >1 ng/m3 (Nessel et al. 1991). In another study, no TCDDs were found in emissions from

hazardous waste or municipal waste incinerators; the levels of PeCDD found in the emissions from

municipal waste incinerators were three orders of magnitude higher than from hazardous waste

incinerators (Oppelt 1991). Fly ash from a municipal incinerator and from coal-fired power plants was

analyzed to study the CDD congener distributions typical of combustion samples (Czuczwa and Hites

1984). OCDD was the most abundant CDD in all fly ash samples. Coal fly ash samples differed

significantly from municipal incinerator fly ash samples. Although some CDDs were detected in coal fly

ash, no TCDDs or PeCDDs were detected. CDDs were present in much lower concentrations in fly ash

from coal-fired power plants than in fly ash from a municipal incinerator. The levels of OCDD in the coal

fly ash samples (2.2 ppb and 3.8 ppb) were at least 100 times lower than those found in the municipal

incinerator fly ash (400 ppb). No isomers of TCDD were detected in municipal incinerator fly ash samples

with a detection limit of 100 ppt (Czuczwa and Hites 1984).

CDDs have been detected in chimney soot samples from various home heating systems using unleaded

heating oil, coal, and wood in Germany (Thoma 1988). A Canadian study of wood-burning stoves

detected only OCDD in particulates from the stack emissions (Wang et al. 1983). Open-air burning of

PCP-treated wood produced levels of CDDs ranging from 2 ppb (TCDD) to 187 ppb (OCDD) (Chiu et al.

1983). Combustion of untreated wood also produces CDDs (TCDD, PeCDD, HxCDD, HpCDD, OCDD)

(Clement et al. 1985). Samples of bottom ash and chimney ash from 2 wood-burning stoves, 1 open

fireplace, and outdoor open-air burning had detectable levels of CDDs ranging from 0.3 to 33 ppb. For

each homologous class, the total concentrations ranged from not detectable to 11 ppb. Detection limits

were equal to 10 ppt for TCDD and PeCDD and 50 ppt for HxCDD, HpCDD, and OCDD. The open-air

burning ash produced the highest total CDD concentration of 33 ppb, with HpCDD (11 ppb) and OCDD

(10 ppb) being the most abundant (Clement et al. 1985).

Fires involving capacitors or transformers containing chlorobenzene and PCBs are also sources of CDDs

and CDFs. For example, in the transformer fire in the New York State Office Building in Binghamton,

NY, TCDD, PeCDD, HxCDD, HpCDD, and OCDD were found in soot samples at levels ranging from

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early 1970s for dust control on roads, parking lots, horse arenas, and other sites around Missouri (Freeman

et al. 1986). The herbicide 2,4,5-T produced commercially prior to 1965 contained up to 30 mg/kg (ppm)

or more 2,3,7,8-TCDD (IARC 1977). The level of 2,3,7,8-TCDD in commercial 2,4,5-T was reduced to

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2,3,7,8-TCDD has been detected in air samples collected at 9 of the 91 NPL hazardous waste sites where it

has been detected in some environmental media (HazDat 1998). Total CDDs have been detected in air

samples collected at 10 of the 126 NPL sites where they have been detected in some environmental media.

Total TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD have been detected in air samples at 10, 3, 3, 3,

and 1 sites of the 105, 34, 43, 49, and 53 sites, respectively, where they have been detected in some

environmental media (see Table 5-1).

5.2.2 Water

CDDs can enter water by a number of different mechanisms including urban runoff, combined sewer

overflows (CSOs), and direct discharge by industrial facilities and publicly-owned treatment works

(POTWs); deposition of particulates from combustion sources, runoff and drift from the use of

chlorophenol-based pesticides; and leaching from chlorophenol-containing waste sites (Huntley et al.

1997; Muir et al. 1986a; Periera et al. 1985; Shear et al. 1996). Direct application or drift of 2,4,5-T or

Silvex into water has also resulted in release of TCDD to surface water (Norris 1981); however, the

contribution of CDDs from pesticide drift is now negligible since most CDD-containing pesticides have

been banned. The migration of chemical wastes containing CDDs from disposal sites has resulted in

contamination of surface water and groundwater (HazDat 1998).

CDDs/CDFs, specifically 2,3,7,8-TCDD and 2,3,7,8-TCDF, are also present in effluent and sludges from

pulp and paper mills that employ the bleached kraft process (Clement et al. 1989; EPA 1991b; Swanson et

al. 1988). 2,3,7,8-TCDD was detected in 7 of 9 bleached pulps at concentrations ranging from not

detected (

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document for the guidelines and standards being proposed for this industry (EPA 1993a). This

development document presents estimates of annual discharges of two congeners, 2,3,7,8-TCDD and

2,3,7,8-TCDF in effluents (from wastewater treatment systems) from this industry as of January 1993.

The joint EPA/paper industry study of 104 pulp and paper mills provides an estimate of the release of

2,3,7,8-TCDD and 2,3,7,8-TCDF in bleached pulp, waste water sludge, and waste water effluent from the

U.S. pulp and paper industry as of mid-to-late 1988 (EPA 1990d). This was a time in the industry’s

history when only limited use of pulping and bleaching technologies and operating practices that

demonstrated potential to reduce the formation of TCDDs and TCDFs had been implemented. In this

study 2,3,7,8-TCDD was detected at 90 and 56% of the kraft and sulfite mills, respectively, that were

surveyed, and no mill was found to be free of 2,3,7,8-TCDD/TCDF. For bleached pulp, the mean

2,3,7,8-TCDD concentration was 7.5 ppt (maximum 56 ppt) for kraft hardwoods, 12 ppt (maximum

116 ppt) for kraft softwoods, 7.1 ppt (maximum 15 ppt) for sulfite hardwoods, and 3.5 ppt (maximum

3.5 ppt) for sulfite softwoods. Mean waste water effluent concentrations of 2,3,7,8-TCDD were 0.076 ppt

for kraft mills (maximum 0.64 ppt) and 0.013 ppt (maximum 0.023 ppt) for sulfite mills. Waste water

sludges contained mean 2,3,7,8-TCDD concentrations of 101 ppt for kraft mills (maximum 1,390 ppt) and

13 ppt (maximum 58 ppt) for sulfite mills. Furthermore, for all kraft mills, about 38% of the

2,3,7,8-TCDD was partitioned to pulps, 33% to waste water sludges, and 29% to waste water effluents.

The NCASI (1993) report found that

CDDs 399


declined by a factor of about 10 from those cited in the 104 Mill Study (EPA 1990d). Overall, NCASI

(1993) reports a 90% reduction in TEQs generated by pulp and paper mills from 1988 to 1992 for all

2,3,7,8-TCDDs and 2,3,7,8-TCDFs.

2,3,7,8-TCDD has been detected in surface water and groundwater samples collected at 9 and 15 sites of

the 91 NPL hazardous waste sites where it has been detected in some environmental media (HazDat 1998).

Total CDDs have been detected in surface and groundwater samples collected at 14 and 32 of the 126 NPL

sites where they have been detected in some environmental media. Total TCDDs, PeCDDs, HxCDDs,

HpCDDs, and OCDD have been detected in surface water samples collected at 10, 1, 4, 4, and 6 sites and

in groundwater samples collected at 21, 3, 10, 14, and 16 of the 105, 34, 43, 49, and 53 NPL sites,

respectively, where these homologues have been detected in some environmental media (see Table 5-1).

5.2.3 Soil

Historically, CDDs have been deposited onto soil through pesticide applications and disposal of CDD-

contaminated industrial wastes, and via land application of paper mill sludges (EPA 1991b). Currently,

however, atmospheric fall-out of CDD-laden particulates and gases appears to be the predominant source

of CDDs to soil (Hutzinger et al. 1985).

The commercial production of trichlorophenol, as well as various derivative products such as 2,4,5-T and

other biocides, has yielded large quantities of waste products containing substantial concentrations of

CDDs. Extensive contamination of the environment with 2,3,7,8-TCDD occurred in Missouri in the early

1970s as a result of the spraying of horse arenas, roads, and parking lots with mixtures of used oil and

chemical waste (Tiernan et al. 1985). The chemical waste, formed during the manufacture of 2,4,5-TCP

and then used to make hexachlorophene, contained several hundred ppm of 2,3,7,8-TCDD (Tiernan et al.

1985). Several thousand gallons of this waste were dispersed over a sizable area of southwestern and

eastern Missouri during the 1970s. Concentrations of 2,3,7,8-TCDD in soil samples from selected

contaminated sites throughout Missouri ranged from 30 to 1,750 ppb (Tiernan et al. 1985). Concentrations

of 2,3,7,8-TCDD in soil samples from Times Beach, Missouri, which had been heavily contaminated,

ranged from 4.4 to 317 ppb (Tiernan et al. 1985).

In Seveso, Italy, an explosion occurred during the production of 2,4,5-T and a cloud of toxic material

including 2,3,7,8-TCDD was released (Cerlisi et al. 1989; MMWR 1988; Mocarelli et al. 1991). Debris

CDDs 400


from the cloud covered an area of approximately 700 acres (2.8 km2). The total amount of 2,3,7,8-TCDD

released during the accident was estimated to be 1.3 kg. Soil samples from this industrial accident were

measured in three areas: Zone A, the most contaminated zone where residents were evacuated; Zone B, the

moderately contaminated area where residents were advised not to eat locally raised produce; and Zone R,

where 2,3,7,8-TCDD contamination in soil was lowest of the three areas. Mean soil concentrations in

these 3 areas were: 230 µg/m2 (maximum 5,477 µg/m2) in Zone A, 3 µg/m2 (maximum 43.9 µg/m2) in

Zone B, and 0.9 µg/m2 (maximum 9.7 µg/m2) in Zone R (MMWR 1988).

The migration of chemical waste containing CDDs from disposal sites has also resulted in environmental

contamination of sediment. For example, at Love Canal in Niagara Falls, New York, where an estimated

200 tons of 2,4,5-TCP production waste were disposed of during the 1940s and early 1950s,

2,3,7,8-TCDD was detected at high concentrations (up to several hundred ppb) in storm sewer sediments

(Smith et al. 1983; Tiernan et al. 1985).

2,3,7,8-TCDD has been detected in soil and sediment samples collected at 61 and 17 sites of the 91 NPL

hazardous waste sites where it has been detected in some environmental media (HazDat 1998). Total

CDDs have been detected in soil and sediment samples collected at 94 and 31 of the 126 NPL sites where

they have been detected in some environmental media. Total TCDDs, PeCDDs, HxCDDs, HpCDDs, and

OCDD have been detected in soil samples at 71, 21, 29, 34, and 38 sites and in sediment samples at 22, 7,

10, 9, and 13 sites of the 105, 34, 43, 49, and 53 NPL sites, respectively, where these homologues have

been detected in some environmental media (see Table 5-1).

5.3 ENVIRONMENTAL FATE

Combustion generated CDDs may be transported long distances (as vapors or associated with particulates)

in the atmosphere (Czuczwa and Hites 1986a, 1986b; Tysklind et al. 1993). They may eventually be

deposited on soils, surface waters, or plant vegetation as a result of dry or wet deposition. CDDs

(primarily MCDD, DCDD, TrCDD) will slowly volatilize from the water column, while the more highly

chlorinated CDDs will adsorb to suspended particulate material in the water column and be transported to

the sediment (Fletcher and McKay 1993; Muir et al. 1992). CDDs deposited on soils will strongly adsorb

to organic matter. CDDs are unlikely to leach to underlying groundwater but may enter the atmosphere on

soil dust particles or enter surface waters on soil particles in surface runoff. Low water solubilities and

high lipophilicity indicate that CDDs will bioconcentrate in aquatic organisms, although as a result of their

CDDs 401


binding to suspended organic matter the actual uptake by such organisms may be less than predicted. This

is also true of uptake and bioconcentration by plants, although foliar deposition and adherence may be

significant.

5.3.1 Transport and Partitioning

Combustion processes appear to have contributed to the ubiquity of CDDs in the environment (Hites and

Harless 1991; Tysklind et al. 1993). CDDs have relatively long residence times in the atmosphere, and

combustion-generated CDDs associated with particulates can become distributed over large areas

(Tysklind et al. 1993). During transport in the atmosphere, CDDs are partitioned between the vapor phase

and particle-bound phase (Hites and Harless 1991). However, because of the very low vapor pressure of

CDDs, the amount present in the vapor phase generally is negligible as compared to the amount adsorbed

to particulates (Paustenbach et al. 1991). The two environmental factors controlling the phase in which the

congener is found are the vapor pressure and the atmospheric temperature (Hites and Harless 1991).

Congeners with vapor pressure 10-4 mm Hg will exist primarily in the vapor phase. Those chemicals

with vapor pressures between these values can be found in both the vapor phase and associated with

particulates (Eisenreich et al. 1981). With a reported vapor pressure ranging from 7.4x10-10 to

3.4x10-5 mm Hg, 2,3,7,8-TCDD falls into the intermediate category.

The detection of CDDs in sediments from Siskiwit Lake, Isle Royale, suggests that CDDs can be

transported great distances in air (Czuczwa and Hites 1986a, 1986b). Because this lake is landlocked on a

wilderness island in Lake Superior, the only way that CDDs could reach these sediments is by atmospheric

fall-out (i.e., by wet and dry deposition). Similar amounts of CDDs were also found in Lake Huron and

Lake Michigan sediments, which indicates that atmospheric transport is a source of CDDs found on these

Great Lake sites (Czuczwa and Hites 1986a, 1986b; Hutzinger et al. 1985). Atmospheric deposition of

TCDD to Lake Erie may contribute up to 2% of the annual input of TCDD to the lake (Kelly et al. 1991).

Through pattern analysis of herring gull monitoring data, Hebert et al. (1994) provided evidence that the

sources of CDDs in Great Lakes food chains were mainly atmospheric, with the exception of

2,3,7,8-TCDD in Lake Ontario, and several CDDs in Saginaw Bay in Lake Huron where point sources

were implicated.

CDDs 402


CDDs are physically removed from the atmosphere via wet deposition (scavenging by precipitation),

particle dry deposition (gravitational settling of particles), and gas-phase dry deposition (sorption of CDDs

in the vapor phase onto plant surfaces) (Rippen and Wesp 1993; Welschpausch et al. 1995). Precipitation

(rain, sleet, snow) is very effective in removing particle-bound CDDs from the atmosphere (Hites and

Harless 1991; Koester and Hites 1992). Table 5-2 summarizes the average ppt scavenging ratios and

percentage of washout due to particulates for congener groups of both CDDs and CDFs collected at two

sites in Indiana. The scavenging ratio is the ratio of the concentration of the congener group in rain to the

atmospheric concentration of the congener group and is a measure of the effectiveness of rain in removing

the congener groups from the atmosphere. Table 5-2 also summarizes the percentages of the congener

groups scavenged as particles in rain rather than as dissolved solutes in rain. Total rain scavenging ratios

ranged from 10,000 to 150,000; HpCDDs and OCDD (the congeners most strongly associated with

particulates) were the congeners scavenged most efficiently (Hites and Harless 1991; Koester and Hites

1992).

Environmental fate modeling of CDDs requires knowledge of a number of fundamental physical and

chemical parameters, such as water solubility, vapor pressure, Henry's law constant, octanol-water

partition coefficient (Kow), and organic carbon partition coefficient (Koc). CDDs are a class of high

molecular weight, highly hydrophobic compounds. Although the class contains 8 homologues (congener

groups) and 75 congeners, solubility values are available for only a handful of these congeners (Doucette

and Andren 1988). CDDs have very low water solubilities, with solubility decreasing with increasing

chlorine substitutions (Doucette and Andren 1988). The water solubility of 2,3,7,8-TCDD ranges from

7.9x10-6 to 33.2x10-4 mg/L (Shiu et al. 1988). See Table 3-2 for the water solubilities for specific

congeners. Water solubilities at 25 EC for the congener groups have been estimated as follows: MCDD,

0.278–0.417 mg/L; DCDD, 3.75x10-3–1.67x10-2 mg/L; TrCDD, 4.75x10-3–8.41x10-3; TCDD, 7.9x10-6 to

6.3x10-4 mg/L; PeCDD, 1.18x10-4 mg/L; HxCDD, 4.42x10-6 mg/L; HpCDD, 2.4x10-6–1.9x10-3 mg/L; and

OCDD, 0.1x10-9–7.4x10-8 mg/L (ASTER 1995; Doucette and Andren 1988; HSDB 1997; McCrady and

Maggard 1993; Shiu et al. 1988).

CDDs generally exhibit very low vapor pressures, with the tendency of decreasing vapor pressure with

increasing chlorine substitution (Friesen et al. 1985; Rordorf 1986, 1989). At 25 EC, the vapor pressure of

2,3,7,8-TCDD ranges from 7.4x10-10 to 3.4x10-5 mm Hg (HSDB 1997; Rordorf 1989). See Table 3-2 for

the vapor pressures of specific congener groups. Vapor pressures at 25 EC for the other congener groups

have been estimated as follows: MCDD, 9.0x10-5–1.3x10-4 mm Hg; DCDD, 9.0x10-7–2.9x10-6 mm Hg;

CDDs 404


6.46x10-8–7.5x10-7; TCDD, 7.4x10-10–4.0x10-3 mm Hg; PeCDD, 6.6x10-10 mm Hg; HxCDD, 3.8x10-11 mm

Hg; HpCDD, 5.6x10-12–7.4x10-8 mm Hg; and OCDD, 8.25x10-13–1.68x10-12 mm Hg (HSDB 1997;

McCrady and Maggard 1993; Rordorf 1989; Shiu et al. 1988). CDDs can be found in both the vapor and

particle-bound phases (Eitzer and Hites 1989a; Hites and Harless 1991), with the low vapor pressure of

OCDD resulting in its enrichment in the particulate phase in the atmosphere. When this particulate matter

is deposited on water, OCDD-enriched sediments will result (Eitzer 1993). The less chlorinated CDD

congeners (TCDD and PeCDD) occur in greater proportion in the vapor and dissolved phases of air and

rain, whereas the more chlorinated congeners (HpCDD and OCDD) are associated with the particulate-

bound phases (EPA 1991d). Data from one study of CDDs in the ambient atmosphere of Bloomington,

IN, found that vapor-to-particle ratios for individual CDDs ranged from 0.01 to 30 and were dependent on

the ambient temperature and the compound's vapor pressure (Eitzer and Hites 1989b). Since the less-

chlorinated CDDs have higher vapor pressures, they are found to a greater extent in the vapor phase (Eitzer

and Hites 1989a). As air moves, photodegradation of the vapor-phase CDDs occurs and they are lost more

readily than the particulate-bound CDDs. Vapor-phase CDDs are not likely to be removed from the

atmosphere by wet or dry deposition (Atkinson 1991), although this is a primary removal process for

particulate-bound CDDs. Wet or dry deposition could result in greater concentrations of the more

chlorinated CDDs reaching soil or water surfaces and eventually sediment (EPA 1991d). All CDDs are

found to some extent in both the vapor phase and bound to particulates. At warmer temperatures (28 EC),

CDDs, particularly the MCDDs, DCDDs, TrCDDs, and TCDDs will have a greater tendency to exist in the

vapor phase. At cooler temperatures (16–20 EC and

CDDs 405


volatilization from the water column is not expected to be a very significant loss process for the TCDD

through OCDD congeners as compared to adsorption to particulates. In general, the Henry's law constants

decrease with increasing chlorine number as a result of the decrease in vapor pressure and water solubility

(Shiu et al. 1988). Volatilization half-lives for 2,3,7,8-TCDD were calculated for ponds and lakes

(32 days) and for rivers (16 days) (Podoll et al. 1986). The primary removal mechanism for CDDs from

the water column is sedimentation, with 70–80% of the CDDs being associated with the particulate phase

(Muir et al. 1992). The remainder was associated with dissolved organic substances. CDDs bound to

sediment particles may be resuspended in the water column if the sediments are disturbed. This could

increase both the transport and availability of the CDDs for uptake by aquatic biota (Fletcher and McKay

1993).

Generally, CDDs are characterized by low vapor pressure, low aqueous solubility, and high hydro

phobicity, suggesting that these compounds strongly adsorb to soil and that their vertical mobility in the

terrestrial environment is low (Eduljee 1987b). In general, higher chlorinated CDDs also volatilize more

slowly from soil and water surfaces than do lower chlorinated ones (Hutzinger et al. 1985). Nash and

Beall (1980) reported that only 12% of 2,3,7,8-TCDD applied to bluegrass turf as a component of

emulsifiable Silvex volatilized over a 9-month period. Because CDDs (particularly the more highly

chlorinated PCDD, HxCDD, HpCDD, and OCDD) strongly adhere to soil and exhibit low solubility in

water, leaching of CDDs would be unlikely if water were the only transporting medium. Instead, wind and

erosion can cause the mixing and transport of CDD-contaminated soil. As a result of erosion, surface soil

contaminated with CDDs is either blown away by wind or washed via surface water runoff into rivers,

lakes, and streams, with burial in the sediments being the predominant fate of CDDs sorbed to soil

(Hutzinger et al. 1985).

Adsorption is an important process affecting transport of hydrophobic compounds such as CDDs. The

organic carbon fraction of the soil is believed to be the most important factor governing the degree of

adsorption of hydrophobic organic contaminants. CDDs adsorb more strongly to soils with a higher organic

carbon content than to soils with low organic carbon content (Yousefi and Walters 1987). Because of their

very low water solubilities and vapor pressures, CDDs found below the surface soil (top few mm) are

strongly adsorbed and show little vertical migration, particularly in soil with high organic carbon content

(Yanders et al. 1989). Vertical movement of CDDs in soil may result from the saturation of sorption sites

of the soil matrix, migration of organic solvents, or human or animal activity (Hutzinger et al. 1985).

Adsorption/desorption of 2,3,7,8-TCDD in contaminated soils was studied by Des Rosiers (1986). Soil

samples were taken from an abandoned 2,4,5-T manufacturing facility and a scrap metal yard in New

CDDs 406


Jersey and from horse arenas, roadways, and residential property in Missouri. Historically, these samples

were contaminated with either chemical residues or waste oils containing 2,3,7,8-TCDD. Mean log organic

carbon partition coefficient (Koc) values ranged from 7.39 to 7.58 (Des Rosiers 1986). This Koc range

indicates that 2,3,7,8-TCDD is immobile in soil (Swann et al. 1983). However, the mobility of

2,3,7,8-TCDD in soil will increase if organic co-solvents that can solubilize 2,3,7,8-TCDD are present in

the soil (Podoll et al. 1986). This situation might occur at a hazardous waste site. In one study, only 1.5%

of the CDDs applied to soil surfaces had leached to a depth of 2.5 cm below the soil surface after

15 months. Leaching of the CDDs through the soil was primarily associated with carriers such as petroleum

oil (Orazio et al. 1992).

Most CDDs entering surface waters are associated with particulate matter (dry deposition of atmospheric

particles) and eroded soil particulates contaminated with CDDs (Hallett and Brooksbank 1986). In the

aquatic environment, significant partitioning of CDDs from the water column to sediment and suspended

particulate organic matter may occur. Dissolved CDDs will partition to suspended solids and dissolved

organic matter (detritus, humic substances) and are likely to remain sorbed once in the aquatic environment.

From suspended sediment and water data collected from the Niagara River on the New York-Canada

border, it was found that CDDs were strongly associated with suspended sediment (Hallett and Brooksbank

1986). Concentrations of total TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD in raw water ranged

from below detection limits to 3.6 pg/L (3.6 ppq), while the concentration of these same homologue groups

in suspended sediments ranged from below detected limits to 228 pg/g (ppt) (Hallett and Brooksbank 1986).

The more highly chlorinated congeners (HxCDD, HpCDD, and OCDD) predominated in both water and

suspended sediment samples.

A model has been developed to describe the vertical transport of low-volatility organic chemicals in soil

(Freeman and Schroy 1986). The model was used to make predictions on the transport of 2,3,7,8-TCDD at

the Eglin Air Force Base Agent Orange biodegradation test plots (Freeman and Schroy 1986). Trenches

10 cm deep were dug in the soil, and Agent Orange containing 40 ppb of 2,3,7,8-TCDD was applied to the

trench bottom. The model predicted a vertical movement of 2,3,7,8-TCDD, buried in 1972, through the soil

column. Soil-column-profile data confirm the vertical movement of 2,3,7,8-TCDD from core samples taken

in 1984 (Freeman and Schroy 1986). The 2,3,7,8-TCDD in the Eglin Air Force Base biodegradation plots

moved through the entire 10 cm of the soil column in 12 years (Freeman and Schroy 1986). The rates of

migration and loss of 2,3,7,8-TCDD in contaminated soil were studied under natural conditions in

experimental plots at the Dioxin Research Facility, Times Beach, Missouri (Yanders et al. 1989). The

CDDs 407


TCDD concentration profiles of sample cores taken at Times Beach in 1988 (mean range 78–160 ppb) were

virtually the same as those in cores taken in 1984 (mean range 76–162 ppb). The results show that little

movement and essentially no loss due to volatilization of 2,3,7,8-TCDD had occurred in the experimental

plots in the four years since the Dioxin Research Facility was established (Yanders et al. 1989).

CDDs are characterized by low water solubilities and high lipophilicities. Kow values range from 104 to 1012

for MCDD through OCDD, with Kow values increasing relative to increasing chlorination (Table 3-2).

Because of these physicochemical properties, CDDs are expected to adsorb to bedded and suspended

sediments and to bioaccumulate in aquatic organisms.

The bioconcentration factor (BCF) is the ratio of the concentration of CDDs in an organism over the

concentration of CDDs in water. The BCF values for CDDs can be estimated from their Kow values, and a

number of regression equations are available for this purpose (Bysshe 1990). Experimentally measured

BCFs for selected CDD congeners in various aquatic species are summarized in Table 5-3. Measurements

of the bioconcentration of CDDs tend to increase with the degree of chlorination up to TCDDs, and then

decrease as chlorination continues to increase up to the OCDD congener (Loonen et al. 1993). The more

highly chlorinated congeners, such as OCDD, appear to have the lowest bioconcentration potential either

because they are less bioavailable because of their rapid adsorption to sediment particles (Servos et al.

1989a, 1989b) or because their large molecule size may interfere with transport across biological

membranes (Bruggeman et al. 1984; Muir et al. 1986a, 1986b).

The hydrophobic nature of CDDs, combined with their great affinity for organic carbon, suggests that a

major proportion of CDDs in the aquatic environment is sorbed to organic matter and sediment. Because

only a minute fraction of CDDs are dissolved in the natural environment, bioconcentration is not the

primary route of exposure for most aquatic organisms. Whereas the term bioconcentration is defined as the

uptake of a chemical from water only, the term bioaccumulation refers to the combined uptake of a

chemical from both dietary sources (e.g., food) and water. A bioaccumulation factor (BAF) that includes

the ingestion route of uptake can be calculated based on fish uptake from water, food, and sediment

(Sherman et al. 1992).

The primary route of exposure to CDD congeners for lower trophic organisms (e.g., phytoplankton and

various aquatic invertebrates) is uptake from the water column or from interstitial water (between sediment

CDDs 410


particles). Certain benthic organisms accumulate highly lipophilic compounds (e.g., PCBs and

CDDs/CDFs) from water at the water/sediment interface (the concentration of a lipophilic compound is

generally higher at this interface than in the water column) and via intake of phytoplankton, zooplankton,

and suspended particulate materials that contain higher concentrations of these chemicals than the

surrounding water (Porte and Albaiges 1993; Pruell et al. 1993; Secor et al. 1993). For the higher trophic

level organisms, such as foraging fish, predaceous fish, and piscivorous wildlife, the predominant route of

exposure is via food chain transfer, with negligible contributions from CDDs in water and sediment (Muir

and Yarechewski 1988). Exposure through direct consumption of CDD-contaminated sediment and detritus

may occur in some bottom-feeding species such as carp and white suckers (Kuehl et al. 1987a, 1987b;

Servos et al. 1989a, 1989b). Under natural conditions, in which a high proportion of these hydrophobic

CDD compounds are sorbed to suspended and dissolved organic matter, direct uptake of these CDDs from

water is not expected to be substantial (Muir et al. 1986a, 1986b). The estimated BCFs in such cases may

not be a good indicator of the experimental bioaccumulation measured in the field. Another reason for the

difference between estimated BCFs and experimentally measured bioaccumulation values is the ability of

some aquatic organisms to metabolize and eliminate specific CDD congeners from their bodies and thereby

change the congener profile pattern in their tissues.

Preferential bioconcentration and bioaccumulation of 2,3,7,8-TCDD and other 2,3,7,8-substituted CDDs

by aquatic organisms have been reported (Branson et al. 1985; Kuehl et al. 1985, 1987a, 1987b, 1987c;

Opperhuizen 1986; Paustenbach et al. 1992). In water-only exposure studies, BCF values for fish exposed

to 2,3,7,8-TCDD ranged from 37,900 to 128,000 (Cook et al. 1991; Mehrle et al. 1988). Much lower BCF

values ranging from 1,400 to 5,840 and 34 to 2,226 have been reported for fish exposed to 1,3,6,8,-TCDD

and OCDD, respectively (Muir et al. 1986a, 1986b). These BCF values are approximately two orders of

magnitude less than would be predicted using the Kow values. Similarly, the lower BCFs for HpCDD in

fathead minnows and OCDD in rainbow trout fry relative to the other CDDs tested resulted from lower

uptake efficiencies from water. Elimination half-lives for TCDDs and PeCDDs were similar and rapid,

averaging about 2.6 days in trout fry and 3 days in minnows. Elimination half-lives for HxCDD and

HpCDD were longer, averaging about 16 days in rainbow trout and 20 days in fathead minnows (Muir et

al. 1986b). The results of these studies also indicate that BCFs of the higher chlorinated CDDs (HxCDD,

HpCDD, OCDD) from water are much lower than would be predicted based on their Kow values. Servos et

al. (1989a, 1989b) also noted that the BCF values were less than predicted based on the Kow values, and

these authors suggest that BCFs reported in the literature may underestimate the true BCF, unless the BCFs

were calculated using truly dissolved CDD concentrations in the water column rather than

CDDs 411


total dissolved concentrations, which would include complexes with large molecules of dissolved organic

carbon.

BCF values measured in fish exposed to both water and sediment were much lower than equivalent

exposures to water only and ranged from 2,500 to 5,800 (Adams et al. 1986; Cook et al. 1991; Tsushimoto

et al. 1982) (Table 5-3). Loonen et al. (1993) also reported that bioaccumulation of CDDs was reduced in

the presence of sediment and that the effects of sediment increased with increasing hydrophobicity (degree

of chlorination) of the congeners. BCFs were reduced by 15–82% for various CDD/CDF congeners, with

the greatest reduction associated with OCDD.

The bioavailability of CDDs/CDFs from municipal incinerator fly ash and sediment to freshwater fish has

been studied in experimental situations. Like the BCF and BAF values, the biota-sediment-accumulation

factor (BASF) (ratio of contaminant concentration in the organism normalized to lipid content to the

concentration in fly ash or sediment, normalized to organic carbon content) generally decreased with an

increasing degree of chlorination (Kuehl et al. 1985, 1987b, 1987c). The BASF values for benthic

(bottom-dwelling) fish (e.g., carp, catfish) are generally higher than for those pelagic (water column)

species (e.g., bass, trout, sunfish) because of the higher lipid content and increased exposure to

contaminated sediments for the benthic species (Paustenbach et al. 1992).

Several authors have studied the disposition and metabolism of CDDs in fish. Studies on the disposition of

2,3,7,8-TCDD in rainbow trout and yellow perch indicate that fatty tissues (visceral fat, carcass, skin, and

pyloric caeca) typically contain the bulk of 2,3,7,8-TCDD (78–90%) with only a small percentage (2–5%)

associated with the skeletal muscle (Kleeman et al. 1986a, 1986b). For other congeners, such as

1,3,6,8-TCDD and OCDD, the greatest proportion of the total body burden is concentrated in the bile, with

lesser concentrations in liver > caeca > kidney > spleen > skin > muscle (Muir et al. 1986a, 1986b).

Differences in the distribution among various species may be a function of the exposure pathway (i.e.,

dietary versus water uptake) and differences in metabolic breakdown rates. For example, both the parent

compound and metabolites of 2,3,7,8-TCDD and 1,3,6,8-TCDD were present in the bile of fish exposed

under laboratory conditions (Branson et al. 1985; Muir et al. 1986a, 1986b). Kleeman et al. (1986b)

reported the presence of several polar metabolites in the gall bladder of yellow perch exposed to a single

dose of 14 C- 2,3,7,8-TCDD. One week later, the gall bladder, skin, skeletal muscle, and kidneys were

removed. In contrast to liver, muscle, and kidney where the parent compound accounted for 96–99% of the

extractable 14 C, the gall bladder contained almost entirely 2,3,7,8-TCDD metabolites, at least one of which

CDDs 412


was a glucuronide conjugate. Although the metabolic breakdown was slow, it is clear that CDDs can be

transformed by fish to polar metabolites that are subsequently excreted in the bile.

Freshwater aquatic invertebrates have been shown to bioaccumulate CDDs/CDFs through water, sediment,

and food pathways (Isensee 1978; Muir et al. 1983; Yockim et al. 1978). The range in experimentally

determined BCF values for freshwater invertebrates is presented in Table 5-3. As discussed previously,

exposure to CDDs from sediment and water containing dissolved organic material markedly decreases the

BCF values, especially for the more highly chlorinated CDDs. Sediment-dwelling organisms (e.g.,

Chironomous sp. larvae and Hexagenia sp. nymphs), stoneflies, and other predaceous nymphs showed poor

accumulation of OCDD in comparison to 1,3,6,8-TCDD (Muir et al. 1983). The lower bioaccumulation of

OCDD was attributed to greater adsorption of the OCDD onto sediment particles and organic matter, and

the reduced uptake across biological membranes due to large molecular size. The potential ingestion of

sediments during burrowing activities by sediment-dwelling insects was believed to result in greater tissue

concentrations of CDDs than those observed for predaceous insects. It is also possible that predaceous

insects may metabolize 1,3,6,8-TCDD more effectively, leading to a greater rate of elimination. Sediment-

dwelling organisms are important food sources for fish and other predaceous insects; consequently, if rapid

elimination of 1,3,6,8-TCDD and low accumulation of OCDD occur in the natural environment,

bioaccumulation of these congeners in trophically higher-level organisms may not be significant (Muir et al.

1983).

Marine invertebrates have also shown an ability to bioaccumulate CDDs/CDFs to varying degrees in their

tissues (Brown et al. 1994; Cai et al. 1994; Conacher et al. 1993; Hauge et al. 1994; Rappe et al. 1991),

although no information on BCF values was found in the literature. Interestingly, several investigators have

reported that shellfish species (crustaceans and molluscs) are better indicators of CDD/CDF contaminant

levels than fish because their tissues contain larger numbers and higher residues of CDD/CDF congeners in

addition to the 2,3,7,8-TCDD congeners and other 2,3,7,8-substituted congeners that are selectively

accumulated in fish species (Brown et al. 1994; Conacher et al. 1993; Rappe et al. 1991). This is in contrast

to what is observed in fish and fish-eating birds, in which there is selective retention of congeners with the

2,3,7,8-substitution positions occupied, which may be due to an increased ability to metabolize and

eliminate non-2,3,7,8-substituted CDD/CDF congeners (Brown et al. 1994; Rappe et al. 1991). The use of

shellfish species as target organisms in CDD/CDF-monitoring studies is recommended as these species

provide a better overall representation of both the magnitude and congener-specific nature of the

environmental contamination (Petreas et al. 1992). Conacher et al. (1993) present an example where

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use of a shellfish species provides a much higher estimate of exposure to CDDs/CDFs as well as to total

CDD equivalent toxicity (TEQs) than use of a fish species. This difference in congener bioaccumulation

profiles between fish and shellfish species is a result of the ability of fish to metabolize CDDs/CDFs. Both

the parent congeners and metabolites of 2,3,7,8-TCDD and 1,3,6,8-TCDD were present in the bile of fish

exposed under laboratory conditions (Branson et al. 1985; Muir et al. 1986a). Kleeman et al. (1986a,

1986b) reported the presence of several polar metabolites, including glucuronide conjugates, in various fish

exposed to 2,3,7,8-TCDD. Despite the slowness of the metabolic breakdown processes, it is clear that

CDDs can be transformed within fish to polar metabolites that are subsequently excreted with the bile. It

does not appear from the results obtained in studies conducted to date that shellfish species have the same

ability to metabolize and eliminate non-2,3,7,8-substituted CDDs/CDFs (Brown et al. 1994; Cai et al.

1994).

It is apparent from the available data regarding the substantial bioaccumulation potential of CDDs/CDFs in

aquatic organisms (particularly the 2,3,7,8-substituted congeners) as well as data on the extent of

contamination of fish and shellfish in various freshwater and marine waterways, that ingestion of

contaminated fish and shellfish is an important exposure pathway for CDDs/CDFs in humans.

CDDs have been found to accumulate in both surface and rooted aquatic vegetation, with BCF values

ranging from 208 to 2,083 (Table 5-3) (Isensee 1978; Tsushimoto et al. 1982; Yockim et al. 1978). Corbet

et al. (1983) reported that a rooted plant species (Potemageton pectimatus) and a surface-dwelling

duckweed (Lemna sp.) accumulated concentrations of 1,3,6,8-TCDD of 280 and 105 ng/g (dry weight),

respectively, following exposure to water containing 1,000 ng/L (ppt). The maximum concentrations were

observed 8 days post-application and represented 6% of the total TCDD applied. These results are similar

to those reported by Tsushimoto et al. (1982) in an outdoor pond study, in which a maximum bioaccumu

lation of 2,3,7,8-TCDD in the pond weeds Elodea nuttali and Ceratophyllon demersum equivalent to a BCF

of 130 occurred after 5 days of exposure. In both studies, the tissue concentrations reached equilibrium in

approximately 20 days and remained constant until the end of the experiment (approximately 58 and

170 days, respectively). These experimental data indicate that CDDs can accumulation in aquatic plant

species through waterborne exposure.

Like many fish, several species of fish-eating birds have shown the ability for preferential bioaccumulation

of 2,3,7,8-TCDD and other 2,3,7,8-substituted CDDs and TCDFs. Jones et al. (1994) monitored TEQ

values for 2,3,7,8-TCDD in double-crested cormorants from three of the Great Lakes: Superior, Michigan,

CDDs 414


and Huron. Biomagnification factors (BMF, the ratio of the concentration of TCDD-equivalents in bird

eggs to concentrations in forage fish) were found to range from 11.7 to 56.8 (mean, 31.3). In another study,

all of the CDDs and CDFs detected in double-crested cormorant and Caspian tern eggs were 2,3,7,8

substituted (Yamashita et al. 1992). Concentrations of 2,3,7,8-TCDD, 1,2,3,7,8-PeCDD, 1,2,3,4,7,8

HXCDD, 1,2,3,6,7,8-HXCDD, 1,2,3,7,8,9-HXCDD, 1,2,3,4,6,7,8-HpCDD, and OCDD ranged from 5.3 to

20, 3.2 to 9.4, 10 to 20, 3.6 to 11, and 7.8 to 16 pg TEQ/g, respectively, for double-crested cormorant eggs,

and 8.2 to 22, 3.3 to 6.4, 8.7 to 17, 2.4 to 6.0, and 9.7 to 21 pg TEQ/g, respectively, for Caspian tern eggs.

This same pattern was also reported to occur in California peregrine falcons and their eggs (Jarman et al.

1993). For this species, mean concentrations of 2,3,7,8-TCDD, 1,2,3,7,8-PeCDD, 1,2,3,4,7,8-HxCDD,

1,2,3,6,7,8-HxCDD, 1,2,3,7,8,9-HxCDD, 1,2,3,4,6,7,8-HpCDD and OCDD in eggs were 5.7, 11, 2, 11, 1.3,

3.8, and 5.3, respectively. Fish-eating birds are exposed to CDDs primarily through their diet. A rapid

decline in contaminant levels in eggs of fish-eating birds, therefore, reflects a rapid decrease in contaminant

levels of their prey. This has been shown to occur in Great blue heron chicks in British Columbia

(Sanderson et al. 1994) in areas where CDD/CDF levels in pulp and paper mill effluents decreased

substantially within a few years. The Great blue heron chicks also showed an increased hepatic microsomal

ethoxyresorufin O-deethylase (EROD) activity in the areas of highest contamination. This indicates that the

induction of cytochrome P-450 1A1 has occurred, and that the Ah-receptor-mediated process, by which

2,3,7,8-TCDD and related chemicals exert their toxicities, has been activated.

Ankley et al. (1993) studied the uptake of persistent polychlorinated hydrocarbons by four avian species at

upper trophic levels of two aquatic food chains. Concentration of 2,3,7,8-TCDD toxic equivalents (TEQs)

were evaluated in Forster’s tern and common tern chicks and in tree-swallow and red-winged-blackbird

nestlings from several areas in the watershed. Young birds accumulated small concentrations of

2,3,7,8-TCDD and several other 2,3,7,8-substituted CDDs and CDFs, including 1,2,3,6,7,8-HxCDD,

2,3,7,8-TCDF, 1,2,3,6,7,8-HxCDF, 1,2,3,4,6,7,8-HpCDF, 1,2,3,7,8-PeCDD, 1,2,3,4,6,7,8-HpCDD, and

OCDD. The general trend in concentrations of CDDs from the greatest to least was Forster’s tern

common tern > tree swallow > red-winged blackbird. The similarity in concentrations between the two tern

species is expected given that they are both piscivores and their similar life histories and the close proximity

of the two colonies. The greater concentrations in the tree swallows than in the red-winged blackbirds were

somewhat unexpected given the presumed similarity of the diets (both species are insectivores). The

authors suspect that the red-winged blackbirds foraged more on relatively uncontaminated upland food

sources than the tree swallows, which fed primarily on chironomids emerging from the bay.

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2,3,7,8-TCDD is generally considered to be bioavailable to terrestrial birds primarily through ingestion of

TCDD-laden food items and soil particles (Nosek et al. 1992). These authors, using H3TCDD-administered

suspensions in various environmental matrices, found that 30% of the dose absorbed from suspensions of

earthworms, 33% absorbed from soil suspensions, 41% absorbed from suspensions of paper mill sludge

solids, and 58% absorbed from a suspension of crickets. These authors also reported that the percentage of

the cumulative TCDD dose translocated to an individual egg was 1.1% for the first 15 eggs laid and that the

percentage was not affected by the order in which the eggs were laid. Assuming an adult female could lay

30 eggs, 35% of the hen TCDD body burden could be translocated to all eggs laid. Results of these studies

suggest that TCDD can be orally bioavailable from earthworms and crickets, important dietary sources for

this species and other terrestrial species, as well as from nonfood items such as orally ingested soil and

paper mill sludge solids.

For terrestrial mammals, the BCF value is the quotient of the concentration of CDD in the tissues divided

by the concentration in food (Geyer et al. 1986a, 1986b). BCF values for 2,3,7,8-TCDD were calculated in

the liver and/or fat of rats, cows, and monkeys (Geyer et al. 1986a; Kociba et al. 1978a). BCF values

ranged from 10.9 to 24.5 in liver tissue and from 3.7 to 24.5 in fat tissue of rats fed 2,200, 210, or 22 ng/kg

of 2,3,7,8-TCDD in their diet for 2 years (Geyer et al. 1986a; Kociba et al. 1978a). The BCF value

calculated for this rat study, increased as the concentration in the animals’ food decreased. In a cattle-

feeding study, 24 ng 2,3,7,8-TCDD in the diet was fed to cows for 28 days after which time the BCF of

2,3,7,8-TCDD in the liver was 0.7 and in the fat was 3.5. Using a linear one compartment model, Geyer et

al. (1986a) calculated that a steady state would be reached in 499 days and that the cattle fatty tissue would

contain 594 ng/kg. The calculated BCF value for 2,3,7,8-TCDD would then be 24.8 (Geyer et al. 1986a;

Jensen et al. 1981). This value is in good agreement with the BCF of 24.5 calculated for rats that received

22 ng TCDD/kg in their diet for years. This is a much higher BCF than has been reported by Fries and

Paustenbach (1990). After 4 years of chronic exposure to 25 ng/kg 2,3,7,8-TCDD in their diet, the

calculated BCF in fatty tissue of monkeys ranged from 24 to 40 (Geyer et al. 1986a). Using the

2,3,7,8-TCDD concentration in human adipose tissue (10.7 ppt whole weight) and in food

(0.052–0.103 ng/kg), the calculated BCF is between 104 and 206 on a whole-weight basis, or between

115 and 229 on a lipid basis (90% lipid) (Geyer et al. 1986a). Using a pharmacokinetics model, the

calculated BCF value is 153 (Geyer et al. 1986a). The authors further point out that the calculated BCFs for

2,3,7,8-TCDD in human adipose tissue are of the same order of magnitude as those calculated for PCBs,

DDT, and hexachlorobenzene which are also persistent compounds with comparable lipophilicity

(n-octanol/water partition coefficients). Based on this BCF range, 2,3,7,8-TCDD was ranked as having a

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high bioconcentration potential in human adipose tissue (Geyer et al. 1986b). The half-life in humans was

estimated to be approximately 7 years (Pirkle et al. 1989).

The primary mechanisms by which CDDs enter terrestrial food chains are by atmospheric wet and dry

deposition of vapor-phase and particulate-bound chemicals (McCrady and Maggard 1993). Uptake of

CDDs from soils by vegetables and other plants may occur (Schroll and Scheunert 1993). Accumulation of

CDDs on vegetation may involve both of these mechanisms. Since 2,3,7,8-TCDD is lipophilic, adsorbs

strongly to soil, and is not very soluble in water, root uptake and translocation to upper plant parts is only a

minor source of vegetative contamination (Travis and Hattemer-Frey 1987) except perhaps for plant species

belonging to the Cucurbitaceas (e.g., zucchini and pumpkin). For zucchini and pumpkin plants, root uptake

of CDD/CDFs and subsequent translocation to the shoots and into the fruits is a main contamination

pathway (Hulster et al. 1994). Hulster and Marschner (1993) reported that CDD levels in foliage were not

related to CDD levels in soil. The contamination of plant foliage via atmospheric deposition is a more

important contamination mechanism than root uptake and translocation to plant foliage (McCrady et al.

1990). Welschpausch et al. (1995) determined that dry deposition was the main pathway of uptake in grass

of CDDs/CDFs from the atmosphere. Particles

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Maize (corn) and bean cultivations grown in soils spiked with 22–1,066 ppt 2,3,7,8-TCDD showed

2,3,7,8-TCDD concentrations in roots ranging from 16 to 1,278 ppt for maize and from 37 to 1,807 for

beans (Fachetti et al. 1986). The soil-grown crops did not show a significant increase of 2,3,7,8-TCDD in

above-ground parts, either as a function of time or with increasing concentration of the pollutant in the soil

(Fachetti et al. 1986).

Uptake of 14C-labeled OCDD was studied in a closed, aerated-soil plant system for 7 days after application

of the OCDD to soil (Schroll et al. 1994). The BCF (concentration of 14C equivalent to the OCDD in plant

dry matter divided by 14C-labeled OCDD in dry soil) was 0.742 in carrot root and 0.085 in carrot shoots

grown on OCDD-contaminated soil as compared to a BCF of not determinable and 0.084 in the control

carrot root and shoots, respectively. There was no transport of 14C-labeled OCDD between the roots and

shoots or vice versa. The residues in roots were due only to root uptake from the soil; those in shoots were

due only to foliar uptake from the air.

Muller et al. (1993) studied transfer pathways of CDD/CDFs to fruit. These authors found that homologue

patterns of CDDs/CDFs in soil were different from those in both apples and pears grown in the contam

inated soil. Concentrations of CDDs/CDFs ranged from 1 to 4 ng/kg (fresh weight) and were 4–8 times

higher in the peel than in the pulp. These authors suggest that airborne CDDs/CDFs are a major source of

contamination of fruits grown in contaminated soil. Muller et al. (1994) conducted field studies of CDD

transfer pathways from soil to several edible plant varieties (carrots, lettuce, and peas). Plants were grown

in soil with 5 ng TEQ/kg or total CDD/CDF concentrations of 363 ng/kg dry weight (control plots) and

56 ng TEQ/kg or total CDD/CDF concentrations of 3,223 ng/kg dry weight on the contaminated plots.

CDD/CDF concentrations in carrot peels were three times higher on the contaminated plots than on the

control plots. This was the result of a 10-fold increase in the CDD/CDF levels in the carrot peel.

CDD/CDF concentrations in lettuce (17.7 and 21.1 ng/kg dry weight) and in peas (7.1 ng/kg dry weight)

were not any higher when grown on the contaminated plot as compared to the control plots and were much

lower than concentrations in the carrots (47.3 and 47.5 ng/kg, dry weight). This indicates that the

CDD/CDFs in the lettuce and peas from both plots were of atmospheric origin. The CDD/CDF homologue

pattern in the contaminated soil showed OCDFs and HpCDFs were the two most prevalent congeners, while

the CDD/CDF homologue pattern from the peel of carrots grown on the contaminated plots contained

TCDF, PeCDF, and HxCDF. Levels of TCDD were the lowest of all CDD/CDF homologues in both

contaminated soils and carrot peels. The homologue profile in lettuce samples was largely dominated by

lower chlorinated CDFs (TCDF and PeCDF) and higher chlorinated CDDs (HpCDD and OCDD), a

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profile often found in samples of atmospheric deposition (Eitzer and Hites 1989a, 1989b). The lowest

CDD/CDF levels of this study were found in peas with pea pods showing higher levels than seeds. The

homologue profiles was dominated by lower chlorinated CDFs and higher chlorinated CDDs similar to the

profile found in lettuce.

Since most of the CDDs released into the atmosphere settle onto water and soil surfaces, foliar deposition is

the major route of vegetative contamination (Travis and Hattemer-Frey 1987). The translocation of foliar

applied 2,3,7,8-TCDD has been studied (Kearney et al. 1971). Labeled 2,3,7,8-TCDD was applied to the

center leaflet of the first trifoliate leaf of 3-week-old soybean plants and the first leaf blade of 12-day-old

oat plants. The compound was applied in an aqueous surfactant solution to enhance leaf adsorption and to

keep the water-insoluble TCDD in solution. Plants were harvested 2, 7, 14, and 21 days after treatment,

dissected into treated and untreated parts, and analyzed. 2,3,7,8-TCDD was not translocated from the

treated leaf to other plant parts. Very little 2,3,7,8-TCDD was lost from soybean leaves, while a gradual

loss (38% in 21 days) did occur from oat leaves (Kearney et al. 1971). The authors considered

volatilization to be a possible mechanism for removal of 2,3,7,8-TCDD, but photolysis may also have

contributed to the loss.

McCrady and Maggard (1993) measured the uptake and elimination mechanisms for 2,3,7,8-TCDD applied

to grass foliage in a closed-laboratory system using [3H]TCDD. The [3H]2,3,7,8-TCDD was injected into

the chamber as a vapor originating from a [3H]2,3,7,8-TCDD generator. The total recovered radioactivity

was 74%. Plant foliage accounted for 59% and the air and other chamber components accounted for 6 and

9%, respectively. This indicated that plant foliage was a major sink for [3H]2,3,7,8-TCDD vapor. Less than

0.2% was recovered from the soil and associated with root tissues, further verifying an airborne mechanism

of [3H]2,3,7,8-TCDD uptake and negligible translocation. The authors also demonstrated that both

photodegradation and volatilization were primary loss mechanisms for [3H]2,3,7,8-TCDD. The

photodegradation half-life (first-order kinetics) of 2,3,7,8-TCDD sorbed to grass and exposed to natural

sunlight was 44 hours, while the half-life for volatilization of 2,3,7,8-TCDD from grass foliage was

128 hours.

In conclusion, CDDs may be transported long distances in the atmosphere. They eventually may be

deposited on soils or surface water as a result of wet or dry deposition. CDDs will slowly volatilize from

the water column or, more likely, will adsorb to suspended particulate materials in the water column and be

transported to the sediment. CDDs deposited on soils will strongly adsorb to organic matter. They are

CDDs 419


unlikely to leach to underlying groundwater, but may enter the atmosphere on soil or dust particles or enter

surface water in runoff. Low water solubilities and high lipophilicity indicate that CDDs will biocon

centrate in aquatic organisms, although as a result of their binding to suspended organic matter, actual

uptake by these organisms may be less than predicted. This is also true of uptake and bioconcentration by

plants, although foliar deposition and adherence may be significant.

5.3.2 Transformation and Degradation

CDDs belong to a class of highly lipophilic compounds with low water solubility and low chemical

reactivity that are resistant to microbial degradation. The dominant transformation processes affecting their

fate have been shown to be surface photolysis and gas-phase diffusion/volatilization with subsequent

photolysis (Yanders et al. 1989).

5.3.2.1 Air

The primary transformation reaction for CDDs in the atmosphere depends on whether the CDD is in the

vapor or particulate phase. Vapor-phase CDDs are not likely to undergo reactions with atmospheric ozone,

nitrate, or hydroperoxy radicals; however, reactions with hydroxyl radicals may be significant, particularly

for the less-chlorinated congeners (MCDD through TCDD) (Atkinson 1991). Based on the photolysis

lifetimes of CDDs in solution, it is expected that vapor-phase CDDs will also undergo photolysis in the

atmosphere, although reactions with hydroxyl radicals will predominate. For TCDD, the photolytic lifetime

ranges from 1.3 to 7.1 days, depending on the season (faster in summer), whereas the hydroxyl radical

reaction lifetime is estimated to be 2 days (Atkinson 1991). A half-life of 8.3 days was estimated for the

gas-phase reaction of 2,3,7,8-TCDD with photochemically produced hydroxyl radicals in the atmosphere

(Podoll et al. 1986). Using the gas-phase hydroxyl radical reaction rate constant of 1×10-11 cm3-molecule-1

sec-1 and an average 12-hour daytime hydroxyl radical concentration of 1.5×106 molecules cm-3, the

atmospheric lifetimes of CDDs are estimated to range from 0.5 days for MCDD to 9.6 days for OCDD, with

TCDD having a lifetime of 0.8–2 days (Atkinson 1991).

Particulate-bound CDDs are removed by wet or dry deposition with an atmospheric lifetime $10 days

(Atkinson 1991) and, to a lesser extent, by photolysis. Miller et al. (1987) measured photolysis of

2,3,7,8-TCDD sorbed onto small-diameter fly ash particulates suspended in air. The results indicated that

fly ash confers photostability to the adsorbed 2,3,7,8-TCDD. The authors reported little (8%) to no loss of

CDDs 420


2,3,7,8-TCDD on the fly ash samples after 40 hours of illumination in simulated sunlight. Koester and

Hites (1992) studied the photodegradation of CDDs naturally adsorbed to five fly ash samples (two from

coal-fired plants, two from municipal incinerators, and one from a hospital incinerator). Although the

authors reported that CDDs underwent photolysis in solution and on silica gel, no significant degradation

was observed in 11 photodegradation experiments conducted for periods ranging from 2 to 6 days.

The selected tran

Documents

5. POTENTIAL FOR HUMAN EXPOSURE · 5. POTENTIAL FOR HUMAN EXPOSURE 5.1 OVERVIEW Chlorinated dioxins (CDDs) are a family of compounds that includes some extremely toxic and potent